Hydrobiologia DOI 10.1007/s10750-013-1459-z
PRIMARY RESEARCH PAPER
Stream regulation by small dams affects benthic macroinvertebrate communities: from structural changes to functional implications Aingeru Martı´nez • Aitor Larran˜aga Ana Basaguren • Javier Pe´rez • Clara Mendoza-Lera • Jesu´s Pozo
•
Received: 25 November 2011 / Revised: 11 January 2013 / Accepted: 19 January 2013 Ó Springer Science+Business Media Dordrecht 2013
Abstract We studied benthic macroinvertebrate communities upstream and downstream of five small reservoirs (surface release in autumn–winters) (north Spain) to assess the effect of flow regulation on structural and functional characteristics of stream ecosystems. We based our approach on the use of structural metrics (density, biomass, richness and diversity) in combination with two functional diversity indices based on biological and ecological traits: FDPG index, related to species richness, and FDQ, which incorporates evenness across taxa. Although water physicochemical parameters were unaffected by the reservoirs during the study period (autumn– winter), macroinvertebrate metrics were lower below the dams, with detritivores (shredders and collector-
gatherers) being the most affected. The alder leaf breakdown rate estimated by the litter-bag technique was related to the density, biomass, richness, diversity and FDPG index of shredders, compromising the ecosystem functioning. The most plausible origin for the observed differences in macroinvertebrate metrics between upstream and downstream reaches was the change of the flow regime caused by the impoundments at downstream sites, leading to droughts in summer in those naturally permanently flowing streams. The observed functional diversity loss might reduce the chances of the community to override natural or man-induced fluctuations in their environment with possible repercussions on important ecosystem functions and services.
Handling editor: Sonja Stendera
Keywords Headwater streams Small dams Benthic macroinvertebrates Functional diversity Ecosystem functioning
Electronic supplementary material The online version of this article (doi:10.1007/s10750-013-1459-z) contains supplementary material, which is available to authorized users. A. Martı´nez (&) A. Larran˜aga A. Basaguren J. Pe´rez J. Pozo Laboratory of Stream Ecology, Department of Plant Biology and Ecology, University of the Basque Country, P.O. Box 644, 48080 Bilbao, Spain e-mail:
[email protected] C. Mendoza-Lera Department of Freshwater Conservation, Brandenburg University of Technology, Cottbus, Seestraße 45, 15526 Bad Saarow, Germany
Introduction Ecological effects of flow regulation have become an important subject in environmental research (Rosenberg et al., 2000). All human societies have dammed rivers with different objectives: water supply, land irrigation, flood control, industrial use and energy generation. In Spain alone, about 1500 reservoirs are documented, of which 35% can be considered small, i.e. dams of less than 15 m of wall height (Spanish
123
Hydrobiologia
Ministry for Agriculture, Alimentation and the Environment; http://www.magrama.gob.es/). These numbers have historically positioned Spain as one of the nations with more dams in the world (see World Commission on Dams; http://www.dams.org/). The impact created by river regulation through large dams has received much attention (Petts, 1988), particularly in temperate regions. These reservoirs may suffer thermal stratification and, through hypolimnetic release, generate substantial changes in the physicochemistry (temperature, oxygen, nutrients, etc.) of the water in the streams below. These changes are responsible for alterations in the biological communities downstream (Haxton & Findlay, 2008). In contrast to their numerical importance, the ecological effects of small dams (\1 hm3, following Martin & Hanson, 1966) have been much less studied than those caused by large dams or hydroelectric production dams. These small dams are usually older and lack any management plan (Poff & Hart, 2002), leading to an uncontrolled release of epilimnetic water. Since the damming impact on fluvial ecosystems depends not only on the size of the dam but also on its use and management (Gore & Petts, 1989), these small dams might create undocumented effects on macroinvertebrate communities. The Serial Discontinuity Concept (Ward & Stanford, 1983) predicts not only changes on water physicochemistry caused by the flow regulation but also changes in the transport of materials (sediments and organic matter) and, in fact, the whole ecosystem functioning. The effects on downstream fluvial biota altered by water chemistry (Fairchild & Velinsky, 2006) and hydrological regime (Marty et al., 2009) have been frequently reported by studies of benthic macroinvertebrate communities (Rehn, 2008). A majority of these studies have focused on structural components of the communities, such as density, biomass or taxa richness (Maxted et al, 2005; Nichols et al, 2006). However, most structural metrics consider species on an equal basis without considering their biological characteristics and ecological role. A common initial approach to study the community structure from a functional point of view consists of the classification of macroinvertebrates in functional feeding groups (Vallania & Corigliano, 2007; Prı´ncipe, 2010) in order to look for trophic ecosystem alterations. Another way to assess impacts on ecosystems that integrate information about biological and ecological traits has been developed as well (Usseglio-
123
Polatera et al., 2000). Although many studies in aquatic systems have demonstrated the response of biological traits to ecosystem alterations (Dole´dec et al., 2006; Pe´ru & Dole´dec, 2010), studies of the effect of small dams on macroinvertebrate traits are lacking so far. Our study addresses the impact of water regulation by small dams on structural and functional metrics of benthic macroinvertebrates. Our main hypotheses are as follows: (1) density and biomass of macroinvertebrates will be lower below the dam, mainly due to changes in hydrology and habitat structure; (2) taxa richness, diversity and functional diversity will also be affected negatively in downstream sites; and (3) these structural and functional losses will correlate to the processing rate of organic matter, which has been previously shown to reduce in the downstream sites of the same systems (Mendoza-Lera et al., 2012).
Materials and methods Study sites The study was carried out in five low-order (2nd–3rd) streams affected by small (0.14–0.64 hm3) surfacereleasing water supply reservoirs (Artiba, Lekubaso, Lingorta, Regato and Zollo) in the Nerbioi-Ibaizabal drainage basin (Northern Spain, mean latitude 438130 N; mean longitude 3870 W), which flows into the Gulf of Biscay (Atlantic Ocean). The climate is temperate with an average annual air temperature of around 148C and a mean annual precipitation of 1200 mm, without droughts in summer (Fig. 1). Water from the entire drainage basin flows through siliceous rock formations and forested watersheds (native Quercus robur L. forests and plantations of Pinus radiata D. Don and P. pinaster Ait.), with alder (Alnus glutinosa (L) Gaertner) dominating the riparian gallery; other anthropogenic land uses are negligible. The reservoirs do not maintain any ecological flow and, consequently, downstream reaches show a year-round variation in water discharge (see Fig. 1, average water level for a 10-year period is shown for the Artiba Reservoir). Autumn seasonal precipitations completely fill the reservoir and it remains at that level until the beginning of summer when the water level gradually drops, whereas with the beginning of autumn, the
Hydrobiologia
This study coincided with an alder leaf-litter decomposition experiment (litter-bag technique) in the same streams (see Mendoza-Lera et al., 2012). During the time that the decomposition experiment was carried out (71 days from November 2008 to February 2009), water temperature was continuously monitored in all sites (data recorded every hour) with SmartButton temperature data loggers (ACR Systems Inc., Surrey, BC, Canada). Conductivity, pH, dissolved oxygen (WTW multiparametric sensor) and river flow (Martin Marten Z30, Current Meter) were
Fig. 1 Average (1999–2008) dammed volume of water (hm3; line) and monthly rainfall (mm; columns) in the Artiba reservoir
water level starts to rise again until winter, when the reservoir is completely full with water (Fig. 1). During the period when the water level is below the maximum level of the reservoir, downstream sites suffer from droughts (confirmed by the authors for the period of the study) and the water level is only recovered by groundwater and inputs from the catchment below the dam. At each stream, two similar 50-m long reaches were selected: one above (Up), in the main tributary of the reservoir, and another one below the dam (Down) (see Table 1 for characteristics of each site; site-specific details in Mendoza-Lera et al., 2012). Due to the accessibility of the streams, downstream reaches were located at different distances from the dam (Artiba 675 m, Lekubaso 220 m, Lingorta 86 m, Regato 285 m and Zollo 660 m). In all ten reaches, riparian vegetation quality was determined using the Riparian Quality Index (QBR; Munne´ et al., 2003) which is based on four descriptors of the riparian vegetation (total riparian vegetation cover, cover structure, cover quality and channel alterations). Benthic habitat quality was estimated using the IHF index (Pardo et al., 2002), which measures the habitat heterogeneity of the substrate and other physical variables of the stream channel. In addition, streambed substrate particle sizes were visually estimated and three categories roughly corresponding to the Wenworth grain size scale (Allan & Castillo, 2007) were used: percentage of ‘boulders’ ([25 cm), ‘cobbles’ (6–25 cm) and ‘gravel-sand’ (\6 cm).
Table 1 Characterization of reservoirs and studied stream reaches. Range of values for watershed surface, reservoir surface, storage capacity and reservoir average depth are shown for the five reservoirs. Range of altitude, reach slope, width, granulometric composition, IHF and QBR and ranges of siteaverage values for temperature, flow, SRP, DIN, alkalinity, conductivity, pH, O2 saturation, FPOM, FPIM, FPM and CPOM for upstream and downstream sites are also shown (period sampled 19th November–2nd February) Reservoir Watershed surface (ha) Reservoir surface (ha)
259–930 2.3–3.9
Storage capacity (hm3)
0.14–0.64
Average depth (m)
6.1–16.5 Up
Down
Altitude (m)
100–340
60–190
Reach slope (%)
3.7–24.4
3.7–27.4
Width (m)
2.8–7.5
4.9–10.9
Water temperature (8C)
7.2–8.8
7.1–7.8 86–675
-1
Flow (l s )
90–250
SRP (lg P l-1)
13.2–18.3
13.4–16.9
DIN (lg N l-1)
196.2–493.0
179.6–597.4
Alkalinity (meq l-1)
0.4–1.2
0.3–1.2
Conductivity (lS cm-1)
94.1–256.8
98.1–254.2
pH
7.5–8.0
7.5–8.0
% O2 Saturation
101.7–109.3
102.4–111.0
FPOM (mg l-1)
0.4–1.3
0.9–2.5
FPIM (mg l-1)
0.4–5.3
1.0–9.9
FPM (mg l-1)
0.8–6.5
1.9–12.4
CPOM (g m-2)
2.7–20.6
3.3–35.1
Granulometric composition (%) [25 cm
29.10–46.67
37.49–53.30
6–25 cm
29.17–49.17
24.15–34.16
\6 cm
20.00–33.33
19.16–30.88
IHF index
75.0–84.0
63.0–79.0
QBR index
67.5–97.5
55.0–85.0
123
Hydrobiologia
measured at 5 sampling dates (ca. fortnightly). In every field visit, water samples were collected, transported to the laboratory and filtered (0.7-mm pore size glass fibre filter, Whatman GF/F). Filters were oven dried (708C, 72 h) and weighed to obtain fine particulate matter (FPM). After combusting the resultant material (5008C, 12 h) and reweighing it, fine particulate inorganic matter (FPIM) was estimated; fine particulate organic matter (FPOM) was calculated as the difference. Alkalinity was determined on filtered water by titration to an end pH of 4.5 (APHA, 2005). Nitrate concentration was determined by the sodium salicylate method, ammonium by the manual salicylate method, nitrite by the sulphanilamide method and soluble reactive phosphorus (SRP) by the molybdate method (APHA, 2005). Benthic standing stock and macroinvertebrates In January 2009, when the remaining alder leaf mass in the bags of the breakdown experiment was approximately 50%, five benthic samples (Surber 0.09 m2, 0.5-mm mesh size) were taken at each study site. The location for each sample was chosen randomly within the available riffles with no predefinition of the substrate to be sampled, which nonetheless showed to be very similar amongst sites (Table 1). Coarse particulate organic matter (CPOM) was sorted out from macroinvertebrates in the field on an 8-mm sieve, oven dried and combusted in the laboratory to determine the ash free dry matter (AFDM). We chose the 8-mm mesh to sieve CPOM as it enabled the separation of the fauna from the organic matter in the field and as the organic matter within the 1–8-mm interval is usually a small fraction of the total. Macroinvertebrates, preserved in 70% ethanol, were identified mainly according to genus (Oligochaeta to order), counted and sorted into functional feeding groups (FFG; after Tachet et al., 2002; Merritt & Cummins, 2007): shredders, collectors-gatherers, collector-gatherer-scrapers, collectors-filterers, scrapers and predators. Individuals from each FFG were altogether oven dried and weighed for dry mass determination. For each sample, we calculated taxa richness, EPT taxa richness, Shannon diversity index and two functional metrics based on 12 selected biological and ecological traits of macroinvertebrates (maximum body size, life cycle length, number of generations per year, dispersion type, consumed food
123
resource, temperature preference, pH preference, trophic status preference, longitudinal distribution, microhabitat preference, locomotion type and current preference). We considered the other 10 traits defined in Tachet et al. (2002) not relevant for our study because (1) we were studying very similar reaches located in a very narrow geographical area and (2) the variation of those traits in our streams was expected to be negligible. The affinity scores for macroinvertebrate traits were obtained from Tachet et al. (2002) and converted into percentages prior to the calculation of the functional metrics. We computed functional diversity indices because the values obtained combine information about the distribution of modalities within traits together with the distribution of modalities amongst species that can thereafter be compared with simple univariate statistics. The two functional diversity metrics computed were Petchey & Gaston’s FDPG (2002) and Rao’s quadratic entropy (FDQ; Rao, 1982; Botta-Duka´t, 2005). Both indices are based on the pairwise difference matrix between taxa of our communities, for which values of trait modalities are taken into account. FDPG is based on the presence–absence of taxa and measures functional diversity by the summed branch lengths of the dendrogram constructed from differences amongst traits, which blends the diversity caused by species richness, number of functional groups, community composition and species identity (Petchey & Gaston, 2002). On the other hand, FDQ is an index of functional diversity that incorporates both the relative abundances of species and a measure of the pairwise functional differences between species. While FDPG ignores the abundance of species, which might undervalue the functional implications of abundant species (Mason et al., 2005), FDQ is influenced by the most abundant species (Rao, 1982), and thus the loss of key species can go unnoticed by it. The comparison of the results obtained by these two metrics can help to discern whether changes in functional diversity are due to changes in taxa richness or/and in the evenness of the abundance. Indices were computed for the entire macroinvertebrate assemblage as well as for each FFG separately since the reservoirs could differently affect the resources sought by the taxa of each FFG. Statistical analyses For physicochemical variables, FPOM, FPIM, QBR, IHF and alder leaf-litter decomposition rates,
Hydrobiologia
comparisons between upstream and downstream sites were presented in Mendoza-Lera et al. (2012). In the present study, these comparisons were related with the faunal variables. Benthic CPOM, density, biomass, taxa richness, EPT richness, Shannon diversity, FDPG and FDQ for total invertebrates and for each FFG (except EPT) were compared using 2-way mixedmodel ANOVAs (factors: ‘stream’, ‘site location’). Relationships between biological parameters and structural indices, FPOM, FPIM, CPOM and leaflitter breakdown rates were tested by ordinary least square linear regressions. Relationships between QBR and macroinvertebrate metrics were tested for each type of site (Up or Down) separately and for the ten sites together; with this approach, we wanted to see if the relationships observed for the upstream reference sites were followed by the impaired downstream sites. When necessary, data were log transformed to meet normality. All statistical analyses were conducted with R (version 2.11.1; R Development Core Team, 2010).
Results During the study period, the dams did not lead to great changes in water physicochemical characteristics; only a slight but significant lower temperature (0.38C) in downstream sites was found. However, reaches below the dams showed a structural simplification as QBR and IHF values were lower downstream (Mendoza-Lera et al., 2012). Total benthic CPOM and deciduous leaves were similar in both types of sites (ANOVACPOM: F1.44 = 0.113, P = 0.739; ANOVADeciduous leaves: F1.44 = 0.001; P = 0.979). Although we did not find statistical differences in discharge between upstream and downstream sites in the winter sampling campaign, we confirmed that the downstream sites were completely dry in summer. A total of 68 taxa were identified, 57 upstream and 55 downstream (Online Resource 1). The average number of taxa per sample was higher in the upstream sites (17) than in the downstream ones (13) (Table 2, ANOVA: F1.44 = 14.692, P \ 0.001). Streams below dam presented lower taxa richness of shredders (50.6% lower; ANOVA: F1.44 = 26.005, P \ 0.001), collector-gatherers (16.5% lower; ANOVA: F1.44 = 6.661, P = 0.013) and collector-gatherers-scrapers (10.6% lower; ANOVA: F1.44 = 5.277, P = 0.026) than
streams above dam. The Shannon diversity of the entire invertebrate assemblages was lower downstream (Table 2; ANOVA: F1.44 = 4.427, P = 0.041) and so was shredders diversity (ANOVA: F1.44 = 18.123, P \ 0.001), collector-gatherers diversity (ANOVA: F1.44 = 5.804, P = 0.020) and collector-gathererscrapers diversity (ANOVA: F1.44 = 20.488, P \ 0.001). EPT richness was 35.6% lower downstream (ANOVA: F1.44 = 22.619, P \ 0.001). The total macroinvertebrate density was highly variable amongst the ten different reaches, ranging from 82.2 individuals m-2 in Lingorta Down to 2940 individuals m-2 in Artiba Up. On average, it was 43.2% lower below the dams (Table 3, ANOVA: F1.44 = 17.616, P \ 0.001). Communities were dominated by detritivores (shredders and any kind of gatherers), which represented per site average 73.36–92.82% for density, and 48.30–99.92% for biomass. There was a decrease in the density of functional feeding groups within detritivores from Up to Down sites: shredders (ANOVA: F1.44 = 11.25, P = 0.002), collector-gatherers (F1.44 = 15.39, P \ 0.001) and collector-gatherer-scrapers (F1.44 = 5.04, P = 0.030). Nevertheless, biomass data did not confirm the observed density reduction (Table 3), attributable to the presence of a few large individuals of Trichoptera for shredders and of Oligochaeta for gatherers in the downstream sites. From the 68 taxa found, a total of 44 showed a lower average density in the downstream sites, although not all differences were statistically significant. The 18 most abundant taxa (those representing [1% of the total density) comprised collectively 91.43% of the total invertebrate abundance. Amongst them, four detritivores and one predator were significantly and negatively affected by the reservoir: the shredders Leuctra and Protonemura (87.3 and 72.7% less abundant downstream; ANOVA: F1.44 = 23.635, P \ 0.001; F1.44 = 8.183, P = 0.006, respectively), the collectorgatherers Habroleptoides, Heptagenia and Orthocladiinae (31.9, 80.8 and 52.5% lower; ANOVA: F1.44 = 12.779, P \ 0.001; F1.44 = 22.605, P \ 0.001; F1.44 = 4.267, P = 0.045, respectively), and the predator Siphonoperla (98.5% lower; ANOVA: F1.44 = 23.405, P \ 0.001). Regarding to the functional diversity indices, FDPG (based on species richness) values were lower below the dam (ANOVA: F1.44 = 11.375, P = 0.002; Table 4). It was also lower downstream in the case
123
Hydrobiologia Table 2 Richness and diversity of total macroinvertebrates and each functional feeding group (FFG), and EPT index (mean ± SE) for upstream and downstream reaches. Results of
2-way ANOVA are shown. The mean and the SE values are based on untransformed data, whereas the statistical output refers to the analysis of the log-transformed data
Up
Down
df
F
P
Comparison
16.76 ± 2.54
12.60 ± 3.24
1.44
14.692
\0.001
U[D
3.56 ± 0.64
1.76 ± 0.46
1.44
26.005
\0.001
U[D U[D
Richness Total Shredders Collector-Gatherers
5.56 ± 0.39
4.64 ± 0.92
1.44
6.661
0.013
Collector-Filterers
1.04 ± 0.34
0.84 ± 0.31
1.44
0.496
0.485
Scrapers
1.56 ± 0.38
1.16 ± 0.40
1.44
1.849
0.181
Col-Gath-Scrap
2.16 ± 0.37
1.84 ± 0.69
1.44
5.277
0.026
Predators
2.88 ± 1.00
2.24 ± 0.67
1.44
1.091
0.302
Total
2.76 ± 0.26
2.43 ± 0.39
1.44
4.427
0.041
U[D
Shredders Collector-Gatherers
1.22 ± 0.26 1.83 ± 0.10
0.51 ± 0.17 1.51 ± 0.32
1.44 1.44
18.123 5.804
\0.001 0.020
U[D U[D
Collector-Filterers
0.30 ± 0.19
0.16 ± 0.09
1.44
1.582
0.215
Scrapers
0.52 ± 0.21
0.28 ± 0.19
1.44
3.27
Col-Gath-Scrap
0.72 ± 0.19
0.40 ± 0.29
1.44
20.488
\0.001
Predators
1.00 ± 0.39
0.84 ± 0.26
1.44
0.279
0.599
EPT
8.20 ± 1.51
5.28 ± 1.34
1.44
22.619
\0.001
U[D
Diversity
Table 3 Density and biomass of total macroinvertebrates and each FFG (mean ± SE) for upstream and downstream reaches. Results of 2-way ANOVA are shown. The mean and SE values Up
Down
0.077 U[D U[D
are based on untransformed data, whereas the statistical output refers to the analysis of the log-transformed data df
F
P
Comparison
Density (ind m-2) 1840.4 ± 455.7
1044.4 ± 380.1
1.44
17.616
\0.001
U[D
Shredders
543.6 ± 316.5
245.3 ± 134.0
1.44
11.249
0.002
U[D
Collector-Gatherers
820.4 ± 156.3
40.6 ± 118.4
1.44
15.388
\0.001
U[D
Total
Collector-Filterers
38.2 ± 11.1
35.1 ± 17.2
1.44
0.366
0.549
Scrapers
85.3 ± 38.5
85.8 ± 35.7
1.44
0.110
0.742
Col-Gath-Scrap
245.3 ± 91.9
220.9 ± 118.8
1.44
5.041
0.030
Predators
119.6 ± 58.5
62.2 ± 29.5
1.44
1.938
0.171
Total
678.1 ± 150.8
891.8 ± 226.0
1.44
0.899
0.348
Shredders Collector-Gatherers
201.1 ± 146.2 232 ± 44.8
211.3 ± 85.1 369.2 ± 135.8
1.44 1.44
1.354 0.458
0.251 0.502
U[D
Biomass (mg m-2)
Collector-Filterers
28.2 ± 15.7
30.7 ± 19.1
1.44
0.850
0.362
Scrapers
47.8 ± 22.6
71.9 ± 38.7
1.44
0.027
0.870
Col-Gath-Scrap
66.1 ± 25.5
59.1 ± 27.3
1.44
4.048
0.051
102.9 ± 38.9
149.5 ± 81.6
1.44
0.312
0.579
Predators
of shredders (ANOVA: F1.44 = 19.540, P \ 0.001) and collector-gatherers-scrapers (ANOVA: F1.44 = 16.743, P \ 0.001; Table 4). FDQ (evenness
123
included) values for total invertebrates did not show differences between the Up and Down reaches (ANOVA: F1.44 = 0.392, P = 0.535); however, the
Hydrobiologia
FDQ value for shredders was lower downstream (Table 4, ANOVA: F1.44 = 9.233, P = 0.004). No physicochemical variable was related to the estimated benthic macroinvertebrate metrics. Considering resource-consumer relationships, suspended FPOM was not related to collector-filterer variables. On the other hand, a positive relationship was found between shredder biomass and the amount of deciduous leaves (Pearson’s R2 = 0.743, P = 0.001, n = 10). The abundance, biomass, richness and FDPG index of benthic shredders showed a significant positive correlation with alder leaf-litter breakdown rates (Fig. 2). The quality of the riparian forest (QBR index) was the variable showing most correlations with the structural and functional metrics based on the total macroinvertebrate community or shredders, but relationships appeared depending on the type of site (altogether, only Up or only Down sites) considered for the regression analysis. When the ten sites were considered together, seven positive relationships appeared. Total density, taxa richness and FDPG for both total invertebrates and shredders (Fig. 3) were following a linear trend with respect to QBR. Considering only Up sites, six significant positive relationships could be seen: Both density and biomass of both total invertebrates and shredders and FDQ of total
invertebrates (Fig. 3) were related to QBR. Down sites did not show any significant relationship with the QBR index. The relationships between macroinvertebrate metrics and IHF index were significant only in three cases: EPT richness, shredder richness and shredder FDPG (Pearson’s R: R2 = 0.401, 0.448, 0.455, P = 0.049, 0.034, 0.032, respectively, n = 10).
Discussion Although not all the studied variables were affected by the dams, structural and functional aspects of the stream always showed significantly lower values below the dam. Changes in water quality caused by river regulation can directly affect the abundance and diversity of aquatic organisms downstream (Bredenhand & Samways, 2009). Nevertheless, we did not observe any noticeable change of physicochemical parameters between upstream and downstream reaches during the study period (Mendoza-Lera et al., 2012). This can be explained by the overflow water release, the small size of the studied dams and their location in oligotrophic headwaters (Cortes et al., 1998; Prı´ncipe, 2010). Water temperature is one of the most frequently affected variables by river
Table 4 Functional diversity indexes, FDPG and FDQ, of total macroinvertebrates and each FFG (mean ± SE) for upstream and downstream reaches. Results of 2-way ANOVA are shown Up
Down
df
F
P
Comparison
FDPG Total
0.31 ± 0.04
0.24 ± 0.05
1.44
11.375
0.002
U[D
Shredders
0.36 ± 0.06
0.26 ± 0.03
1.44
19.540
\0.001
U[D U[D
Collector-Gatherers
0.44 ± 0.02
0.36 ± 0.07
1.44
5.536
0.023
Collector-Filterers
0.16 ± 0.08
0.12 ± 0.06
1.44
0.878
0.354
Scrapers
0.19 ± 0.09
0.18 ± 0.08
1.44
0.005
0.944
Col-Gath-Scrap Predators
0.47 ± 0.07 0.18 ± 0.07
0.24 ± 0.15 0.16 ± 0.05
1.44 1.44
16.743 0.582
\0.001 0.449
U[D
FDQ Total
0.41 ± 0.03
0.39 ± 0.07
1.44
0.392
0.535
Shredders
0.19 ± 0.05
0.09 ± 0.04
1.44
9.233
0.004
Collector-Gatherers
0.46 ± 0.03
0.39 ± 0.09
1.44
2.921
0.094
Collector-Filterers
0.06 ± 0.04
0.03 ± 0.03
1.44
0.319
0.575
Scrapers
0.08 ± 0.04
0.09 ± 0.06
1.44
0.055
0.815
Col-Gath-Scrap
0.22 ± 0.09
0.19 ± 0.12
1.44
0.297
0.589
Predators
0.21 ± 0.07
0.25 ± 0.08
1.44
0.718
0.402
U[D
123
Hydrobiologia Fig. 2 Relationship between structural and functional variables for shredders and alder leaflitter breakdown rate. Filled symbols represent upstream sites and open symbols represent downstream sites. Streams are identified as follows: Artiba (circle), Lekubaso (inverted pyramid), Lingorta (pyramid), Regato (diamond) and Zollo (square). Note the different scales
impoundment (Pozo et al., 1997; Bredenhand & Samways, 2009), and it is a very important factor for the biology and the evolutionary ecology of stream insects (Ward & Stanford, 1982). We reject the idea that minimal changes in water temperature and in other physicochemical parameters can explain the observed differences in invertebrate communities, but state that other factors are more important determinants for invertebrates. Independent of the stream, the CPOM to FPOM ratio suffers a natural decline as the detritus is transported downstream, but impoundments generate steep reductions of the ratio due to coarse detritus entrapment (Ward & Stanford, 1983). This disruption, as Bredenhand & Samways (2009) observed, causes alterations in the functional trophic structure of the macroinvertebrate community below dams leading to an increase of the number of filterers and a decrease of shredders. Nevertheless, in our study, despite the higher amount of FPOM in suspension downstream, we did not observe higher densities of collector-filterers below dams. On the other hand, although dams did not affect the amount of CPOM, the richness and densities of shredders and collectorgatherers were reduced downstream. Our downstream sites suffer a year-round flow fluctuation that approximates to that observed in Mediterranean streams: with flows similar to temperate streams throughout autumn–winter–spring when the
123
reservoirs are filled, and droughts during summer (observed by the authors), since the dams in the present study retain all the water without ensuring any ecological flow downstream. Navarro-Lla´cer et al. (2010) noted that these alterations of the annual flow regime induced a loss in riparian forest quality downstream of impoundments under Mediterranean climate. In low-order forested streams, riparian forests define the functional feeding structure of macroinvertebrate communities (Sandin, 2009) by determining the quality and quantity of the allochthonous resources (Sweeney, 1993). Flow alteration is also able to homogenize the benthic habitat structure (Hart & Finelli, 1999), which can determine the distribution of benthic biota (Sandin, 2009), as many taxa require specific habitat types (see Tachet et al., 2002). Substratum homogenization was observed in our downstream sites (Mendoza-Lera et al., 2012), but it was not related to the expected increase of generalist taxa recorded by Townsend et al. (1997). However, the significant correlations found between QBR and macroinvertebrates indicate that structural changes in the riparian corridor might influence macroinvertebrate communities. The positive and significant relationship between deciduous leaves and shredder biomass is also pointing out the importance of allochthonous inputs for the invertebrates. Nevertheless, the correlation between QBR and
Hydrobiologia b Fig. 3 Relationship between riparian vegetation quality index
(QBR) and structural and functional variables for total macroinvertebrate assemblage and shredders. Filled symbols represent upstream sites and open symbols represent downstream sites. Streams are identified as follows: Artiba (circle), Lekubaso (inverted pyramid), Lingorta (pyramid), Regato (diamond) and Zollo (square). Significant linear regressions are shown for the 5 upstream sites (continuous line) and for the 10 studied sites (dashed line). No significant relationship was observed using only the 5 downstream sites. Note the different scales
macroinvertebrates was only evident in the upstream sites, with no significant correlation if we only consider the downstream sites. This contrasting result between
the upstream and the downstream sites might indicate that the riparian forest is an important driver of macroinvertebrate communities in natural streams (our upstream sites) and that the impact on downstream sites has its origin in the observed flow alteration. Actually, water shortage caused by summer droughts can by itself define the structure of communities (Mun˜oz, 2003). After dry periods, the taxonomic composition depends on the recolonization capacity that varies amongst taxa (Otermin et al., 2002) and notably the scarcity of shredders in comparison with other FFG at these first stages (Boulton, 1991). Instead of temporal changes, some authors have studied streams with different flow regimes. For instance, Sabater et al. (2008) compared a permanent Atlantic stream with an intermittent Mediterranean stream and found lower detritivore and predator biomass in the intermittent stream. We have not observed differences between upstream and downstream sites in terms of biomass, but abundance results are comparable to Sabater’s study with a significantly lower shredder density in reaches below the dam and also lower densities of Leuctra and Protonemura (shredders), Habroleptoides, Heptagenia and Orthocladiinae (collector-gatherers), and Siphonoperla (predator), most of them described as slow colonizers (Gore, 1982; McArthur & Barnes, 1985; Otermin et al., 2002). We lack macroinvertebrate data for summer months, but the patterns of upstream–downstream differences in macroinvertebrate communities in winter and the visually corroborated droughts in downstream reaches in summer suggest that Up–Down differences might become even higher. Macroinvertebrate assemblages from most of the streams contain many species that are redundant (Lawton, 1991) in the sense that ecosystem functions can proceed even when some of the taxa are absent (Wallace et al., 1986). Nevertheless, our trait-based results of functional diversity indices support that
123
Hydrobiologia
alterations in the structural metrics of the community are accompanied by a change at the functional level as shown by the FDPG index which was lower downstream and which is strongly correlated with taxa richness (Petchey & Gaston, 2002). The effects were, nevertheless, undetected by the FDQ index. This index includes evenness in the computation and is further strongly influenced by abundant taxa (Rao, 1982); hence, changes of functionally singular but rare taxa (amongst them keystone species) can go unnoticed. Although it is difficult to consider any stream macroinvertebrate a keystone species (Paine, 1966; Mills et al., 1993), there are some ecosystem processes in which some taxa seem to be cornerstones (McKie & Cranston, 1998; Woodward et al., 2008). One of the most important processes in low-order forested streams is the coarse organic matter processing, for which shredders are mainly responsible. In our streams, this function was negatively affected by river regulation as we observed a reduction of above 20% in leaf-litter decomposition rates from upstream to downstream sites (Mendoza-Lera et al., 2012). This reduction in the decay could be more easily explained by structural metrics of benthic shredders than by metrics of shredders associated with bags. In fact, we obtained higher R2 values using benthic data than those reported by Mendoza-Lera et al. (2012) using the fauna of litter bags. Leaf-litter processing is just one amongst many other functions that are essential for freshwater ecosystems. The loss of species and the consequent loss of trait diversity would also weaken the capability of the community to satisfactorily develop other functions. Considering stream regulation by small dams, we have observed that the taxa involved in the detritic pathway seem to be the most affected (shredders and gatherers have suffered significant reductions in both functional diversity indexes, FDPG and FDQ). If there are functions, apart from leaf-litter decomposition, in which shredder and collector-gatherer taxa are especially important (e.g. secondary production, recolonization, etc.), regulated headwater streams might also suffer from deficiencies in those other processes. Nevertheless, caution should be observed when stating that small dams weaken the resistance of a given community because of two reasons. Firstly, the functional redundancy amongst taxa (Lawton, 1991) can lead to adaptations of the communities to the loss of species, filling the gaps left by missing taxa by others with similar characteristics.
123
Since density and biomass of shredders were good predictors of the rates, being irrelevant the inter-site taxonomic differences on shredder assemblages, redundancy might be playing for the leaf-litter decomposition. Secondly, we have not observed upstream–downstream differences for the biomass of macroinvertebrates. Biomass correlates stronger than density to other important properties (i.e. metabolism) and functions (i.e. secondary production) of the macroinvertebrate community and the lack of differences in biomass might be understood as a limited impact. To sum up, the regulation by surface release small reservoirs negatively affects density, richness and diversity of macroinvertebrates in the reaches below the dam. The main driver is probably the variability of the flow regime due to the absence of any ecological flow, which creates droughts below the dam in summer. Furthermore, downstream reaches are characterized by loss of riparian forest quality and reduction of benthic habitat heterogeneity. The changes in the trophic structure of the macroinvertebrate communities are strongly related to stream processes, such as leaf-litter decomposition, and are accompanied by a loss of functional diversity, which in turn could alter functions of the system. Acknowledgments This study was funded by the Spanish Ministry of Science and Innovation (project CGL2007-66664C04-01) and by the Basque Government (Research grant IT422-07, IT-302-10). A. Martı´nez was provided with a grant by the Basque Government and J. Pe´rez by the University of the Basque Country. We are grateful to L. L. Federlein and two anonymous referees for the suggestions that improved the manuscript.
References Allan, J. D. & M. M. Castillo, 2007. Stream ecology: structure and function of running waters, 2nd ed. Springer, Dordrecht: 436 pp. APHA (American Public Health Association), 2005. Standard Methods for the Examination of Water and Wastewater, 21st ed. American Public Health Association, American Water Works Association, and Water Environment Federation, Washington, DC: 1368 pp. Botta-Duka´t, Z., 2005. Rao’s quadratic entropy as a measure of functional diversity based on multiple traits. Journal of Vegetation Science 16: 533–540. Boulton, A. J., 1991. Eucalypt leaf decomposition in an intermittent stream in south eastern Australia. Hydrobiologia 211: 123–136.
Hydrobiologia Bredenhand, E. & M. J. Samways, 2009. Impact of a dam on benthic macroinvertebrates in a small river in a biodiversity hotspot: Cape Floristic Region, South Africa. Journal of Insect Conservation 1: 297–307. Cortes, R. M. V., M. T. Ferreira, S. V. Oliveira & F. Godinho, 1998. Contrasting impact of small dams on the macroinvertebrates of two Iberian mountain rivers. Hydrobiologia 389: 51–61. Dole´dec, S., N. Phillips, M. Scarsbrook, R. H. Riley & C. R. Townsend, 2006. Comparison of structural and functional approaches to determining landuse effects on grassland stream invertebrate communities. Journal of the North American Benthological Society 25: 44–60. Fairchild, G. W. & D. J. Velinsky, 2006. Effects of small ponds on stream water chemistry. Lake and Reservoir Management 22: 321–330. Gore, J. A., 1982. Benthic invertebrate colonization: source distance effects on community composition. Hydrobiologia. 94: 183–193. Gore, J. A. & G. A. Petts, 1989. Alternatives in regulated river management. CRC Press, Boca Raton, FL: 344 pp. Hart, D. D. & C. M. Finelli, 1999. Physical-biological coupling in streams: the pervasive effect of flow on benthic organisms. Annual Review of Ecology and Systematics 30: 363–395. Haxton, T. J. & C. S. Findlay, 2008. Meta-analysis of the impacts of water management on aquatic communities. Canadian Journal of Fisheries and Aquatic Sciences 65: 437–447. Lawton, J., 1991. Are species useful? Oikos 62: 3–4. Martin, R. O. R. & R. L. Hanson, 1966. Reservoirs in the United States, Water Supply Paper 1838. U.S. Geological Survey, Washington, DC: 115 pp. Marty, J., K. Smokorowski & M. Power, 2009. The influence of fluctuating ramping rates on the food web of boreal rivers. River Research and Applications 25: 962–974. Mason, N. W. H., D. Mouillot, W. G. Lee & J. B. Wilson, 2005. Functional richness, functional evenness and functional divergence: the primary components of functional diversity. Oikos 111: 112–118. Maxted, J. R., C. H. McCready & M. R. Scarsbrook, 2005. Effects of small ponds on stream water quality and macroinvertebrate communities. New Zealand Journal of Marine and Freshwater Research 39: 1069–1084. McArthur, J. V. & J. R. Barnes, 1985. Patterns of macroinvertebrate colonization in an intermittent rocky mountain stream in Utah. Great Basin Naturalist 45: 117–123. McKie, B. G. L. & P. S. Cranston, 1998. Keystone coleopterans? Colonization by wood-feeding elmids of experimentally immersed woods in south-eastern Australia. Marine and Freshwater Research 49: 79–88. Mendoza-Lera, C., A. Larran˜aga, J. Pe´rez, E. Descals, A. Martı´nez, O. Moya, I. Arostegui & J. Pozo, 2012. Headwater reservoirs weaken terrestrial-aquatic linkage by slowing leaf-litter processing in downstream regulated reaches. River Research and Applications 28: 13–22. Merritt, R. W. & K. W. Cummins, 2007. An introduction to the aquatic insects of North America. Kendall/Hunt Publishing Company, Dubuque: 1158 pp. Mills, L. S., M. E. Soule & D. F. Doak, 1993. The keystonespecies concept in ecology and conservation. BioScience 43: 219–224.
Munne´, A., N. Prat, C. Sola, N. Bonada & M. Rieradevall, 2003. A simple field method for assessing the ecological quality of riparian habitat in rivers and streams: QBR index. Aquatic Conservation: Marine and Freshwater Ecosystems 13: 147–163. Mun˜oz, I., 2003. Macroinvertebrate community structure in an intermittent and a permanent Mediterranean streams (NE Spain). Limnetica 22: 107–116. Navarro-Lla´cer, C., D. Baeza & J. de las Heras, 2010. Assessment of regulated rivers with indices based on macroinvertebrates, fish and riparian forest in the southeast of Spain. Ecological Indicators 10: 935–942. Nichols, S., R. Norris, W. Maher & M. Thoms, 2006. Ecological effects of serial impoundment on the Cotter River, Australia. Hydrobiologia 572: 255–273. Otermin, A., A. Basaguren & J. Pozo, 2002. Re-colonization by the macroinvertebrate community after a drought period in a first-order stream (Agu¨era basin, Northern Spain). Limnetica 21: 117–128. Paine, R. T., 1966. Food web complexity and species diversity. American Naturalist 100: 65–75. ´ lvarez, J. Casas, J. L. Moreno, S. Vivas, N. Bonada, Pardo, I., M. A J. Alba-Tercedor, P. Ja´imez-Cue´llar, G. Moya`, N. Prat, S. Robles, M. L. Sua´rez, M. Toro & M. R. Vidal-Abarca, 2002. El ha´bitat de los rı´os mediterra´neos. Disen˜o de un ´ındice de diversidad de ha´bitat. Limnetica 21: 115–133. Pe´ru, N. & S. Dole´dec, 2010. From compositional to functional biodiversity metrics in bioassessment: a case study using stream macroinvertebrate communities. Ecological Indicators 10: 1025–1036. Petchey, O. L. & J. G. Gaston, 2002. Functional diversity (FD), species richness and community composition. Ecology Letters 5: 402–411. Petts, G. E., 1988. Impounded rivers: perspectives for ecological management. John Wiley & Sons Ltd, West Sussex: 326. Poff, N. L. & D. D. Hart, 2002. How dams vary and why it matters for the emerging science of dam removal. BioScience 52: 659–738. Pozo, J., E. Orive, H. Fraile & A. Basaguren, 1997. Effects of the Cernadilla-Valparaiso reservoir system on the River Tera. Regulated Rivers: Research and Management 13: 57–73. Prı´ncipe, R. E., 2010. Ecological effects of small dams on benthic macroinvertebrate communities of mountain streams (Co´rdoba, Argentina). Annales de LimnologieInternational Journal of Limnology 46: 77–91. R Development Core Team (2010) R: A language and environment for statistical computing. R Foundation for Statistical Computing, Vienna, Austria. ISBN 3-900051-07-0, URL: http://www.R-project.org. Rao, C. R., 1982. Diversity and dissimilarity coefficients: a unified approach. Theoretical Population Biology 21: 24–43. Rehn, A. C., 2008. Benthic macroinvertebrates as indicators of biological condition below hydropower dams on west slope Sierra Nevada streams, California, USA. River Research and Applications 25: 208–228. Rosenberg, D. M., P. McCully & C. M. Pringle, 2000. Globalscale environmental effects of hydrological alterations: introduction. BioScience 50: 746–751. Sabater, S., A. Elosegi, V. Acun˜a, A. Basaguren, I. Mun˜oz & J. Pozo, 2008. Effect of climate on the trophic structure of
123
Hydrobiologia temperate forested streams. A comparison of Mediterranean and Atlantic streams. Science of the Total Environment 390: 475–484. Sandin, L., 2009. The effects of catchment land-use, near-stream vegetation, and river hydromorphology on benthic macro invertebrate communities in a south-Swedish catchment. Fundamental and Applied Limnology 174: 75–87. Sweeney, B. W., 1993. Effects of streamside vegetation on macroinvertebrates communities of White Clay Creek in eastern North America. Proceedings of the Academy of Natural Sciences of Philadelphia 144: 291–340. Tachet, H., P. Richoux, M. Bournaud & P. Usseglio-Polatera, 2002. Inverte´bre´s d’eau douce: syste´matique, biologie et e´cologie. CNRS, Paris: 587 pp. Townsend, C. R., S. Dole´dec & M. R. Scarsbrook, 1997. Species traits in relation to temporal and spatial heterogeneity in streams: a test of habitat template theory. Freshwater Biology 37: 367–387. Usseglio-Polatera, P., M. Bournaud, P. Richoux & H. Tachet, 2000. Biological and ecological traits of benthic freshwater macroinvertebrates: relationships and definition of groups with similar traits. Freshwater Biology 43: 175–205.
123
Vallania, A. & M. C. Corigliano, 2007. The effect of regulation caused by a dam on the distribution of the functional feeding groups of the benthos in the sub basin of the Grande river (San Luis, Argentina). Environmental Monitoring and Assessment 124: 201–209. Wallace, J. B., D. S. Vogel & T. F. Cuffney, 1986. Recovery of a headwater stream from an insecticide-induced community disturbance. Journal of North American Benthological Society 5: 115–126. Ward, J. V. & J. A. Stanford, 1982. Thermal responses in the evolutionary ecology of aquatic insects. Annual Review of Entomology 27: 97–117. Ward, J. V. & J. A. Stanford, 1983. The serial discontinuity concept of lotic ecosystems. In Fontaine, T. D. & S. M. Bartell (eds.), Dynamics of lotic ecosystems. Ann Arbor Science Publishers, Michigan: 29–42. Woodward, G., G. Papantoniou, F. Edwards & R. B. Lauridsen, 2008. Trophic trickles and cascades in a complex food web: impacts of a keystone predator on stream community structure and ecosystem processes. Oikos 117: 683–692.
